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Emissions of Sulfur Trioxide from Coal-Fired Power

Plants

R.K. Srivastava a , C.A. Miller a , C. Erickson b & R. Jambhekar b

a

 Office of Research and Development, National Risk Management Research Laboratory,

Air Pollution Prevention and Control Division , U.S. Environmental Protection Agency ,

Research Triangle Park , North Carolina , USA

b

 Babcock Borsig Power, Inc , Worcester , Massachusetts , USA

Published online: 22 Feb 2012.

To cite this article: R.K. Srivastava , C.A. Miller , C. Erickson & R. Jambhekar (2004) Emissions of Sulfur

Trioxide from Coal-Fired Power Plants, Journal of the Air & Waste Management Association, 54:6, 750-762, DOI:

10.1080/10473289.2004.10470943

To link to this article: http://dx.doi.org/10.1080/10473289.2004.10470943

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Emissions of Sulfur Trioxide from Coal-Fired Power Plants

R.K. Srivastava and C.A. Miller

Office of Research and Development, National Risk Management Research Laboratory, Air

Pollution Prevention and Control Division, U.S. Environmental Protection Agency, Research

Triangle Park, North Carolina

C. Erickson and R. Jambhekar

Babcock Borsig Power, Inc., Worcester, Massachusetts

ABSTRACT

Emissions of sulfur trioxide (SO3) are a key component of

plume opacity and acid deposition. Consequently, these

emissions need to be low enough to not cause opacity

violations and acid deposition. Generally, a small fraction

of sulfur (S) in coal is converted to SO3 in coal-fired

combustion devices such as electric utility boilers. The

emissions of SO3 from such a boiler depend on coal S

content, combustion conditions, flue gas characteristics,

and air pollution devices being used. It is well known that

the catalyst used in the selective catalytic reduction (SCR)

technology for nitrogen oxides control oxidizes a small

fraction of sulfur dioxide in the flue gas to SO3. The extent

of this oxidation depends on the catalyst formulation and

SCR operating conditions. Gas-phase SO3 and sulfuric

acid, on being quenched in plant equipment (e.g., air

preheater and wet scrubber), result in fine acidic mist,

which can cause increased plume opacity and undesirable

emissions. Recently, such effects have been observed at

plants firing high-S coal and equipped with SCR systems

and wet scrubbers. This paper investigates the factors that

affect acidic mist production in coal-fired electric utility

boilers and discusses approaches for mitigating emission

of this mist.

INTRODUCTION

As understanding of the adverse effects of air pollution

has grown, so also has the complexity of coal-fired power

plant design and operation, especially with regard to air

pollution control systems. Control of air pollutant emissions is not only a legal requirement but also is becoming

a financial necessity, as salability of effluents and trading

of emissions increase the direct monetary value of emissions control. The days when one must only consider the

nuisance value of fly ash are long past.1

As plant complexity has increased, so have unexpected consequences of changing segments of the total

chemical process that occurs between fuel preparation

and ultimate emissions. One of the more discernible adverse consequences is the formation and emission of sulfur trioxide (SO3)/sulfuric acid (H2SO4), as highlighted by

the recent and well-publicized experiences of a power

plant in Ohio.2 Although not directly subject to emission

limits, SO3 is important to consider during the design and

operation of coal-fired utility boilers for a number of

environmental and plant performance reasons.

The formation of SO3 will occur during the combustion of sulfur (S)-bearing fuels such as coal and heavy fuel

oils. Virtually all of the SO3 converts to H2SO4 as flue gas

is cooled in the air preheater (APH). Relatively high concentrations of SO3/H2SO4 in the boiler, stack, or plume

can cause adverse impacts to plant equipment and to the

environment. Impacts on plant equipment can include

corrosion, fouling, and plugging and may require additional hardware or changes in operation to minimize SO3/

H2SO4 concentrations and the resulting adverse impacts.

Health and Environmental Effects

Formation of visible H2SO4 droplets depends upon

the concentration and dew point of H2SO4 and the

IMPLICATIONS

Formation and emissions of SO3/H2SO4 can potentially

lead to plant operation-related problems and environmental

concerns. The emissions of SO3 from a boiler depend on a

number of interacting complex factors including coal sulfur

content, combustion conditions, flue gas characteristics,

and air pollution devices being used. Understanding the

paramenters leading to excessive generation of SO3/H2SO4

emissions can assist in selection of practical control approaches. This paper investigates the factors that affect

SO3/H2SO4 emissions from coal-fired electric utility boilers

and discusses approaches for mitigating these emissions.

The paper is expected to be a state-of-the-art reference on

the subject for regulators, utility industry personnel, and

others stakeholders.

TECHNICAL PAPER ISSN 1047-3289 J. Air & Waste Manage. Assoc. 54:750–762

Copyright 2004 Air & Waste Management Association

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concentration of sub-micron particles upon which the

acid can condense. As seen in Figure 1, H2SO4 dew point

is a function of water vapor and H2SO4 concentration and

increases as both these variables increase.3 The curves in

Figure 1 determine the fraction of H2SO4 in the vapor and

condensed phases for temperatures below the dew point.

For example, a 250 °F flue gas with 10% water vapor and

8 ppm H2SO4 will have a vapor-phase acid gas concentration of 2 ppm, with the remaining 6 ppm in the condensed phase.

If temperatures at the stack are low enough and water

vapor and H2SO4 concentrations are high enough, the

condensed-phase concentration of H2SO4 can be at a level

that results in the formation of a visible plume attached to

the stack. Once the gases leave the stack, the rate of

cooling and subsequent H2SO4 condensation competes

with the plume dilution by entrainment of ambient air

into the plume. Even in cases where stack conditions are

such that the H2SO4 is completely in the vapor phase at

the stack exit, a detached plume may still form shortly

downwind as temperatures drop below the dew point.

In most cases, and particularly in the case where

H2SO4, water, and sub-micron solid particles are present,

condensation is the dominant formation mechanism.

The opacity of these plumes is most strongly influenced

by the concentrations of condensing species and submicron fly ash particles present in the stack gases. The size

distribution of the sub-micron fly ash also can have a

noticeable effect on plume opacity at low to moderate

H2SO4 concentrations, because these particles act as condensation sites for the condensing vapor-phase H2SO4.

The relationship between sub-micron particle concentration, H2SO4 concentration, and plume opacity is shown

in Figure 2 for flue gas with a sub-micron particle mode at

0.15 m diameter.4,5 The gray area denotes a typical range

of sub-micron particle concentrations downstream of an

electrostatic precipitator (ESP).6

The curves in Figure 2 illustrate that H2SO4 concentration can have a strong impact on opacity when fine

particle concentrations are present at realistic levels. Because the condensing acid particles can nucleate to form

particles even at very low levels of pre-existing submicron particles, further reduction of sub-micron solid

particles may not significantly reduce plume opacity. As

an example, a unit emitting 16 ppm of vapor-phase

H2SO4 and 5 mg/m3 of sub-micron particles is predicted

to result in a plume opacity of 35%. An opacity of 20%

can be achieved by reducing sub-micron PM concentration to 1.6 mg/m3 (i.e., 70% reduction) or by reducing

H2SO4 to 8 ppm (a 50% reduction). For units with

limited ability to further reduce PM emissions, control of

SO3 (and, subsequently, H2SO4) may be the only viable

option for meeting plume-opacity requirements. Because

essentially all the SO3 is converted to H2SO4 at or below

stack temperatures, Figure 2 also illustrates the impact of

increases in SO3 formation on plume opacity. A unit with

1 mg/m3 of sub-micron particles in the stack gases is

predicted to experience doubling of plume opacity from

5 to 10% when H2SO4 concentrations increase from 5 to

10 ppm.

Data on direct adverse effects on human health are

inconsistent, but studies indicate that aqueous acidic

aerosols at typical and even elevated ambient concentrations generally have minimal effects on symptoms and

mechanical lung function in young healthy adults. However, there are studies that have shown changes in mucociliary clearance and modest bronchoconstriction in asthmatics exposed to elevated concentrations of 400 g/m3

or more.7,8 Of more concern than the effects of exposure

Figure 1. Dew point of H2SO4 as a function of H2SO4 concentration

for different water vapor concentrations.3

Figure 2. Iso-opacity plot calculated for a bimodal distribution of

coarse and sub-micron particles. The sub-micron particle mode is at

0.15 m diameter.5 The gray area represents typical sub-micron mass

concentrations in the stack of a coal-fired utility boiler.6

Srivastava, Miller, Erickson, and Jambhekar

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to typical ambient levels of SO3/H2SO4 aerosol alone is

exposure to this aerosol mixed with other ambient PM

constituents, including elemental carbon and metals.9

In situations where SO3/H2SO4 aerosol combines

with a sinking plume, the ambient concentrations near

the stack can reach significantly higher levels than those

normally experienced. In such cases where meteorological and operating conditions combine to form a nearstack acidic fog, damage to property and vegetation can

occur if the conditions are sustained for an extended

period. Anecdotal evidence of adverse health impacts,

such as burning eyes, sore throats, and headaches, has

also been reported in such cases.2

Impacts on Plant Hardware

The most common problem associated with elevated concentrations of SO3 is low-temperature (300 °F [150 °C])

corrosion.10 Once formed, SO3 reacts easily with the moisture in combustion flue gases to produce H2SO4. Below its

dew point, H2SO4 condenses and will collect in relatively

low-temperature areas of the flue gas path, such as the air

heater, and corrode the contacted metal components. The

dew point varies with H2SO4 and water concentrations in

the flue gas (see Figure 1), but is typically between 200 and

300 °F (95 and 150 °C, respectively), corresponding to low

(8%) moisture, low (0.1 ppm) H2SO4 conditions and high

(15%) moisture, high (20 ppm) H2SO4 conditions, respectively. At higher temperatures (above 1000 °F [540 °C]),

SO3 can cause corrosion in superheaters and reheaters.11

These high-temperature reactions are less common than

the low-temperature condensation and corrosion reactions. In addition to forming corrosive H2SO4, SO3 reacts

to form sulfate (SO4

2) particles, especially ammonium

sulfate [(NH4)2SO4] and ammonium bisulfate (NH4HSO4),

in boilers using ammonia (NH3), which can lead to fouling or plugging of APH passages.12 Regular maintenance is

conducted at plants to address the impacts described. This

maintenance can include monitoring and replacement of

materials, washing of APH passages, and other activities.

In some cases, modification or addition of hardware

may be required to minimize corrosion or plugging. At

plants using wet flue gas desulfurization (FGD) systems,

the reduction in flue gas temperature can result in increased H2SO4 condensation and subsequent corrosion.

In such cases, the flue gas may require reheating to ensure

it remains above the dew point, thereby resulting in increased operation costs.13

REGULATORY OVERVIEW

There are currently no U.S. regulations that directly limit

emissions of SO3/H2SO4 aerosols from utility boilers.

Other regulatory programs may, however, require SO3

control, as discussed later. German emission standards

limit the combined sulfur dioxide (SO2) and SO3 emission

concentrations to a daily average of 50 mg/Nm3 and a

30-min average of 200 mg/Nm3

. There is no separate

emission limit for SO3 alone.14

In the United States, SO3 emissions are included as

part of limits on opacity and particulate matter (PM)

emissions, and H2SO4 is listed as a hazardous air pollutant

(HAP) under Title III of the Clean Air Act Amendments of

1990.15 Under Title III, electric utility generating stations

are exempt from the mass emission limits of 10 t (9.09

tonnes) per year of any single HAP or 25 t (22.7 tonnes)

per year of combined HAPs set for other industrial

sources. Thus, although H2SO4 is a listed HAP and is

emitted in significant quantities by coal-fired utility generating stations,16 there is no regulatory requirement to

control these emissions as HAPs. They must, however, be

reported to the U.S. Environmental Protection Agency

(EPA) under the Toxics Release Inventory (TRI) requirements. The H2SO4 is typically measured and reported as

equivalent SO3.

EPA has defined primary PM as particles that are

either emitted directly as a solid or liquid or are emitted as

a vapor but condense or react upon cooling and dilution

in the ambient air to form solid or liquid PM immediately

after discharge from the stack.17 Secondary PM is composed of particles that form through chemical reactions in

the ambient air well after dilution and condensation have

occurred.18 Under this definition, SO3 or H2SO4 that is in

the vapor phase in the furnace or stack and either reacts to

form SO4

2 particles or condenses to form liquid H2SO4

immediately after leaving the stack is considered part of

primary PM emissions and is therefore subject to PM

emission limits.

In some states or localities, plants may be required to

control SO3 to maintain opacity standards. Even though

opacity measurements in the stack may be within regulatory limits, the formation of a visible plume caused by the

presence of SO3 may result in plume opacities higher than

allowable levels and a need to control SO3 emissions.

Most states have a 20% opacity limit for stationary

sources, including coal-fired utility boilers, and many also

have provisions for higher opacity levels during unit

startup. There are some exceptions to the 20% limit, with

a few states having opacity limits of 40% and at least one

(West Virginia) with a 10% opacity limit.

The formation of a detached plume is indicated when

in-stack opacity measurement values are lower than those

in the downwind plume. While in-stack opacity measurements are made using continuous emission monitors,

measurements of the downwind plume opacity are covered under EPA Method 9.9 The increase in opacity from

in-stack measurements to downwind is often caused by

the condensation of vapors such as H2SO4 and can result

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in regulatory violations and good neighbor policy concerns.

FORMATION OF SO3

Formation in the Boiler

The S in coal has inorganic and organic components. The

inorganic component is predominantly pyrite, which can

exist both as distinct particles (excluded) and bound in

the coal matrix (included). The organic component is part

of the various organic structures present in coal. The

fractions of included and excluded pyrite and organic

component delivered to a furnace are known to be dependent on the fineness to which the coal is ground. During

the combustion of coal, virtually all the S gets oxidized to

gaseous species such as SO2 and SO3, with SO2 being the

principal oxide. Although detailed chemical mechanisms

for oxidation of fuel S are not understood well at present,

it is believed that this oxidation proceeds through rapid

formation of SO2, occurring on a timescale comparable to

that of fuel oxidation reactions. Because SO2 formation is

so rapid, its concentration can be estimated using equilibrium calculations.

The primary reaction that results in SO3 formation in

flames is

SO2  O  M 3 SO3  M (1)

This recombination reaction proceeds rapidly near the

combustion zone in the presence of super-equilibrium

concentrations of oxygen (O) atoms. The consumption of

SO3 near flames occurs primarily via

SO3  HO2 3 HOSO2  O2 (2)

and

SO3  O 3 SO2  O2 (3)

A few other reactions are considered to result in formation

and destruction of SO3 near the flame zone, but the effect

of these on SO3 concentrations is relatively minor. Table

1 summarizes the reactions involving SO3 in the combustion zone.19

To understand SO3 formation, calculations were

made using the S chemistry described previously with gas

compositions and temperature-time history characteristic

of a coal-fired boiler.20 This history is shown in Figure 3.

Initial gas-phase compositions are assumed to result from

complete combustion (with 10 and 30% excess air) of a

Montana sub-bituminous and a Western Kentucky bituminous coal. The sub-bituminous coal is composed of

53.26% (by weight) carbon (C), 3.35% hydrogen (H),

0.87% nitrogen (N), 0.78% S, 9.34% ash, 21.23% water,

and 11.16% O. The bituminous coal contains 69.79% C,

4.79% H, 1.34% N, 2.95% S, 7.47% ash, 5% water, and

8.65% O. The initial compositions resulting from complete combustion of the coals are then equilibrated at

1647 °C (1900 K) to provide the input composition for the

chemical kinetic calculations based on the S chemistry.

Note that the H2-O2-carbon monoxide (CO) subset of the

hydrocarbon mechanism given in ref 21 is used in these

calculations. This subset includes 25 reactions involving

11 species.

Figures 4 and 5 plot SO3 mole fraction and the ratio

SO3 mole fraction/SO2 mole fraction versus time, respectively. Because reaction 1 is exothermic, little SO3 is

formed initially at high temperatures in the furnace. As

the flue gas cools, SO3 is produced, and the rate of production is high in the convective region of the boiler (see

Figures 3 and 4). SO3 production via gas-phase S kinetics

is complete before the flue gas enters the economizer. The

results also reflect that SO3 production increases with coal

Table 1. The reactions involving SO3 in the combustion zone. The rate constants

are listed in the form k  AT  exp(EA /RT ) with units in mol-cm-sec.

Reaction A  EA /R

SO3  H  HOSO  O 2.5e05 2.92 25,300

SO3  O  SO2  O2 2.0e12 0 10,000

SO3  SO  SO2  SO2 1.0e12 0 5000

SO2  O  M  SO3  M 9.2e10 0 1200

SO2  OH  SO3  H 4.9e02 2.69 12,000

HOSO2  SO3  H 1.4e18 2.91 27,600

HOSO2  O  SO3  OH 5.0e12 0 0

HOSO2  OH  SO3  H2O 1.0e12 0 0

HOSO2  O2  SO3  HO2 7.8e11 0 330

Figure 3. Temperature-time history for a coal-fired power plant.20

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S content and furnace excess air level. This is because

increases in the S content of coal or furnace excess air

level result in a corresponding increase in the concentration of SO2 and/or O, which, in turn, results in greater

production of SO3 via reaction 1. Finally, Figure 5 reflects

that SO2 conversion rate (i.e., SO3/SO2 molar ratio) for the

modeled coals ranges between 0.1 and 0.65%, approximately, and is relatively independent of coal type at a

specific excess air level. Because boilers are operated with

between 10 and 30% excess air, a 0.65% conversion of

coal S to SO3 should provide a conservative estimate of

SO3 production in the furnace.

In addition to the SO3 formation in the furnace discussed above, additional formation takes place in the

temperature range of 1100–800 °F (593–427 °C) found in

the economizer region of the boiler. This formation results from oxidation of SO2 via molecular oxygen (O2)

catalyzed by iron oxides present in both ash and tube

surfaces.22 This oxidation mechanism depends on several

operating and design parameters, including SO2 concentration, ash content and composition, convective pass

surface area, gas and tube surface temperature distributions, and excess air level. Because the impact of these

parameters depends on site-specific factors (e.g., cleanliness of tube surfaces), it is difficult to characterize the

extent of SO3 formation caused by catalytic oxidation of

SO2. A laboratory study found that conversion of SO2 to

SO3 in the presence of fly ash ranged between 10 and

27% for temperatures between 500 and 700 °C, and essentially increased linearly with the iron oxide content of

the ash.23 These conversion rates are substantially higher

than those found in coal-fired boilers, for which data

suggest furnace/economizer conversion to be 0.8 to

1.6% for bituminous and 0.05–0.1% for sub-bituminous

coals.24 The above laboratory study results indicate that

temperatures and residence times greater than, and carbon ash contents lower than, those typically found in

coal-fired utility boilers are required to achieve high oxidation rates.

Formation of SO3 in SCR Reactors

The SCR technology is increasingly being used at power

plants to control emissions of oxides of nitrogen (NOx). In

the SCR process, NH3 is injected into the flue gas within a

temperature range of 600–750 °F (315–400 °C), upstream of a catalyst. Subsequently, as the flue gas contacts

the SCR catalyst, NOx, which predominantly is nitric oxide (NO) in combustion devices, is chemically reduced to

molecular nitrogen (N2). In the most commonly used SCR

Figure 4. SO3 produced during coal combustion.

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process layout, known as hot-side SCR, the catalyst is

located between the economizer and the APH.

It is well known that the catalyst used in the SCR

technology oxidizes a small fraction of SO2 in the flue gas

to SO3:25–29

SO2  1⁄2 O2 3 SO3 (4)

The extent of this oxidation depends on the catalyst formulation and SCR operating conditions. Generally, this

oxidation can range from 0.25 to 0.5% of SO2 in bituminous and from 0.75 to 1.25% in low-S sub-bituminous

coal applications. To examine what this oxidation means,

consider an SCR application with an SO2 oxidation guarantee of 0.5%. Also assume that the concentrations of SO2

and SO3 in the flue gas at the inlet of the SCR reactor are

2000 and 20 ppm, respectively. Then, based on oxidation

across the catalyst, the concentrations of SO2 and SO3 at

the exit of the SCR reactor will be 0.995  2000  1990

ppm and 2000  1990  20  30 ppm, respectively. Thus,

the SO3 loading in the flue gas at the exit of the SCR

reactor will be 50% more than that at the inlet. This

example illustrates that the level of oxidation across the

catalyst can have a significant impact on SO3 concentration.

In general, for a given catalyst material experiencing

the same flue gas conditions, the oxidation rate of SO2 to

SO3 (or conversion rate) is inversely proportional to area

velocity (AV), which is simply the ratio of flue gas volumetric flow rate to geometric catalyst surface area. This

implies that the conversion rate is proportional to catalyst

volume (and, hence, geometric surface area) and gas residence time in the catalyst. Thus, the conversion rate, 	,

can be expressed as

	  K/AV (5)

where the constant of proportionality, K, is a function of

catalyst material and design as well as flue gas properties;30 that is,

K  f(catalyst material, catalyst design,

flue gas properties) (6)

Figure 5. SO3 conversion during coal combustion.

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A systematic study of oxidation of SO2 to SO3 over honeycomb SCR catalysts has been reported in ref 31. The

findings of this study are as follows:

(1) The conversion rate depends primarily on the

vanadium content of the catalyst and, therefore,

can be controlled by adjusting this content;

(2) The oxidation reaction is considerably slower

than diffusion of SO2 within the pores of the

catalyst. Therefore, the entire volume of the catalyst is active in oxidation of SO2 to SO3 in contrast to reduction of NOx to N2, which, being

diffusion-limited, occurs mainly at the catalyst

surface. The rate of oxidation is linearly proportional to catalyst wall thickness. Accordingly, reducing the wall thickness should not affect NOx

reduction but should reduce SO2 to SO3 conversion;

(3) The reaction rate is of variable order in SO2 concentration, increases with temperature, is independent of concentrations of O2 and water in

practical applications, is strongly inhibited by

NH3 and is slightly enhanced by NOx.

SO3 DEPLETION OR CONVERSION TO H2SO4

Processes in APHs

Utility boilers use APHs to transfer heat from hot flue gas

exiting the economizer to combustion air flowing into the

boiler. These APHs are available in rotary regenerative and

tubular designs, with the former used more widely. Typically, the flue gas temperature at the APH inlet is between

600 and 700 °F (316 and 371 °C) and 300 °F (149 °C) at

the exit. SO3 is hygroscopic and, therefore, absorbs vaporphase moisture at temperatures above its dew point to

form H2SO4 vapor.22 This process occurs in the APH. The

extent of conversion of SO3 to H2SO4 depends on the

temperature distribution in the APH and flue gas moisture

content. However, virtually all SO3 converts to H2SO4 at

temperatures of 400 °F or less. If local metal temperatures

in the APH flow passages drop below the acid dew point,

some H2SO4 condenses on these surfaces as liquid droplets (aerosol). This rate of condensation is dependent on

the wall temperature and H2SO4 concentration in the flue

gas.

In a regenerative APH, where flow passages are periodically exposed to hot flue gas and relatively cold incoming combustion air, evaporation of condensed H2SO4 occurs on exposure to air. The rate of this evaporation is

dependent on the moisture content of air and metal surface temperature of the APH flow passage.32,33

Results of an extensive field test program indicate

that 40% of the flue gas SO3 present at the regenerative

APH inlet is removed in the APH by the condensationevaporation mechanism discussed above.34 The tubular

APH design does not have surfaces that are periodically

exposed to combustion air; except for small leakages in

welds and seal, no mixing of the flue gas and combustion

air occurs. Therefore, the resultant H2SO4 formed is not

removed by evaporation into the combustion air and

passes directly out of the APH. Modeling of the acid condensation phenomenon in tubular APHs is described in

ref 35.

In addition to the condensation-evaporation of SO3

discussed above, an additional conversion process takes

place in boilers using SCR. In such applications, NH3 is

injected as a reagent in the SCR process. A minor fraction,

2–5 ppmv, of injected NH3 slips past the SCR catalyst and

does not react with NOx. This fraction of NH3, known as

NH3 slip, reacts with SO3 downstream of the SCR reactor

and forms (NH4)2SO4 and NH4HSO4 salts and, thereby,

results in removal of SO3 from the flue gas.12 This salt

formation can be detrimental to the APH performance if

APH passages become plugged and pressure loss across the

APH results in forcing offline washing. The amount and

type of salt formed will depend on the amount of NH3

slip; based on typical concentrations of SO3 and H2O in

flue gas, the amount of NH3 slip is the limiting factor in

salt-formation reactions.

Localized Condensation of H2SO4 in the Duct

between APH and PM Control

Condensation of H2SO4 also can take place in the duct

between the APH and the PM control device.35,36 For

example, because of the rotating heat transfer element

employed in the regenerative APH, gas flow stratification

across the flow cross-section at the APH inlet is enhanced

in the APH. As a result, strong transverse variations in gas

temperature and H2SO4 vapor concentration can exist in

the gas leaving the APH. These variations can lead to

localized condensation of H2SO4. Based on operating conditions, localized condensation also can occur in units

using tubular APHs. For a given combination of flue gas

H2SO4 concentration and moisture content, Figure 1 can

be used to determine if condensation is occurring.

Fly Ash Adsorption and Removal of H2SO4 in PM

Control Equipment

In addition to the processes described, some H2SO4 gets

adsorbed on fly ash in the APH and downstream equipment. The rate of this adsorption depends on the temperature of the flue gas, concentration of H2SO4, and fly ash

properties, in particular alkalinity.23,37 The adsorption

may increase rapidly as the flue gas reaches the cold end

of the APH and may continue in the duct between the

APH and the PM control equipment. The adsorbed H2SO4

gets removed with the fly ash in the PM control device

(ESP or baghouse).

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The firing of sub-bituminous coals, which generally

have S contents on the order of 0.5%, results in fly ash

with a relatively high amount of alkali (20–30% by

weight). Such alkaline ash adsorbs virtually all H2SO4 in

the flue gas.38 In such cases, SO3 injection is required for

improving ESP performance. In contrast, based on fly ash

properties and temperature, the majority of SO3 in flue

gas of a bituminous coal-fired boiler may or may not be

adsorbed in fly ash.30 Finally, hot-side ESPs operate at

high-enough temperatures where little adsorption of

H2SO4 occurs.

As mentioned above, adsorption of H2SO4 on fly ash

depends on the alkalinity of the fly ash and the concentration of H2SO4. Figure 6 illustrates the general adsorption characteristics of fly ash as a function of the molar

ratio of alkali content of fly ash and SO3 concentration at

the inlet of the APH. The alkali content of fly ash is

defined as the molar sum of magnesium oxide (MgO) and

calcium oxide (CaO) in the ash. This figure represents a

correlation of data from Babcock Borsig Power field-testing programs concerning SO3 production and capture in

flue gas systems.

Aerosol Formation in Wet FGD Systems

As the H2SO4 vapor-containing flue gas passes through a

wet FGD system, it is rapidly cooled below the acid dew

point. Because the rate of this cooling is greater than the

rate of absorption of H2SO4 vapor in the scrubber solution, the dew-point crossover results in H2SO4 mist with

sub-micron droplets.30,39 In a wet scrubber, generally, the

mass transfer between this mist and the scrubbing liquid

occurs through inertial impaction, gravitational settling,

Brownian diffusion, diffusiophoresis, and thermophoresis.

For sub-micron mist droplets with diameters less than

0.05 m, Brownian diffusion is known to be the primary

component of mass transfer.40 In practical situations, depending on the relative velocity between the sub-micron

mist droplets and the collecting liquid form (droplets,

wetted walls, liquid sheets, and bubbles), the Brownian

motion-induced mass transfer is often not adequate to

result in high-efficiency capture of sub-micron droplets.

Therefore, generally, larger droplets in the mist can be

removed in the scrubber, but a significant portion of the

sub-micron droplets are not removed and are emitted

from the stack. This explains why wet FGD systems are

relatively inefficient in removing SO3/H2SO4. In general,

50% of the H2SO4 entering the scrubber may be removed in the scrubber.

MEASUREMENT OF SO3

The controlled condensation system (CCS) has been used

since the 1960s for the measurement of SO3 and H2SO4 in

flue gas streams.41,42 The general arrangements of the CCS

and thimble holder are shown in Figures 7 and 8, respectively. The measurement system consists of a quartz-lined

heated probe that draws gas through a quartz thimble for

the removal of PM. The probe and thimble are heated to

avoid condensation of SO3 vapor in the gas sample. The

gases enter a temperature-controlled condenser where the

SO3 is condensed on the wall to be collected and measured after the sample run via a deionized water rinse.

The measurement techniques for SO3 and H2SO4 recently have been under review for accuracy and improvement for TRI reporting. Recently, a review and verification

project has been undertaken by EPRI to qualify the

CCS method and compare field data with available

Figure 6. Fly ash SO3 capture rate downstream of furnace.

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prediction methods used to estimate SO3 emissions for

TRI reporting.37 The EPRI project compared laboratory

data with field test data for a variety of coals with a range

of S and fly ash chemical compositions. The study further

investigated the SO3 removal rates and efficiencies of the

plant equipment used for field verification.

The EPRI project has concluded that alkaline ash will

produce a bias in the SO3 measurement. The bias is a

result of SO3 removal in the thimble holder by the alkaline ash collected on the quartz thimble. This bias is seen

as significant for low-S, high-alkaline ash fuels (Powder

River Basin coals). The CCS was found to have a bias of

20–25% low readings for high S, acidic ash fuels and

greater than 40% for Powder River Basin coals. A possible

solution to the bias is utilized in Europe.43 The test

method uses a system similar to the CCS; however, the

quartz thimble is replaced with a small tubular ESP. The

tubular ESP removes the ash to the sidewalls away from

the gas stream; in contrast, the quartz thimble filters the

flue gas through the collected ash. Furthermore, the European system can distinguish between gaseous and aerosol SO3 using deionized water procedures on various collection plates. The European SO3 method system has been

used extensively in the United States on a variety of coals

with repeatable and reliable results.

MITIGATION OF SO3 EMISSIONS

The mitigation of SO3 has been an active area of research

for many years.13,23,44,45 This research has resulted in the

development, refinement, and implementation of different techniques and methods for the successful mitigation

of SO3 in the flue gases of fossil-fuel–fired boilers. Recently, a guide that summarizes SO3 mitigation options

and their respective success in either full-scale or pilot

testing has been written.10 The mitigation options are

described below.

Alkali Addition into the Furnace

The injection of alkali in the flue gas stream has been a

method of SO3 mitigation for almost 30 years.44 The location in the flue gas stream and the delivery method of

the alkali have been studied and tested depending on

operating and site conditions.

The addition of alkali into the furnace recently has

been proven to be effective at the full scale.30,39 However,

the method of delivery and the effectiveness of the alkali

used varies from site to site depending on specific conditions. The addition of limestone to coal before pulverization for the control of SCR catalyst arsenic poisoning has

been shown to be an effective method of furnace SO3

control. Recent commissioning data46 have shown at least

Figure 7. CCS sampling train.

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a 50% reduction in furnace SO3 from limestone addition.

The injection of alkaline sorbents (Ca- and Mg-based slurries) has been shown to be effective at controlling SO3

emissions of the furnace.39 The slurry injection method

has successfully obtained SO3 furnace conversion reductions of 40–80% but has been found to be sensitive to

injection location and elevation. The effects of the MgO

sorbent slurries on SCR catalyst activity are currently under investigation. To date, no studies have measured the

potential benefit of SCR catalyst arsenic poisoning control

by MgO sorbent injection. The addition or injection of

alkali in the furnace does not influence the conversion

rate of the SCR catalyst.

Alkali Injection after the Furnace

The injection of other alkaline materials after the furnace/

economizer exit has been used for the control of SO3 for

both APH corrosion and stack emissions.10 The primary

sorbents used are compounds such as hydrated lime, limestone, MgO, and sodium carbonate. The selection of a

sorbent for a given site will depend on economic factors,

such as availability and required SO3 removal rates. Alkali

injection has successfully removed between 40 and 90%

of the SO3 in the flue gas at various plants depending on

the injected material and injection rate.10 The sorbents

are introduced into the flue gas as either a dry powder or

mixed with water to form slurry before injection. The

location of injection in the flue gas stream varies. Plants

with APH cold-end corrosion problems may elect to inject

the sorbent before the APH; however, adequate APH

cleaning equipment is required with this configuration.

In general, the common injection location for SO3 control is between the APH and the ESP. The injection of

sorbent before the ESP must, however, consider the effect

on particulate control. The ESP will have higher inlet

mass loading, and the fly ash will have different resistivity

characteristics. It has been reported that dry injection of

hydrated lime has resulted in strong sparking and lower

operating currents in the ESP during pilot-scale testing.10

Ammonia Injection before the ESP

The injection of NH3 after the APH and before the ESP has

been shown to be 90% effective in the removal of SO3 in

full-scale application.10 This method of mitigation results in the formation of (NH4)2SO4 and NH4HSO4 salts in

the ESP, depending on the NH3 and SO3 concentration

Figure 8. CCS thimble holder.

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ratios. The formation of NH4HSO4 is expected when NH3

to SO3 molar ratios are less than 1; this formation tends

to decrease the ESP particle loading caused by fly ash

agglomeration. NH3 injection before the ESP is used for fly

ash conditioning to increase ESP performance caused by

the agglomeration effects. With NH3 to SO3 molar ratios

between 1 and 2, increased formation of (NH4)2SO4 is

expected, with an increase in the particle loading of the

ESP. The injection of NH3 results in the adsorption of NH3

by the fly ash. Because the fly ash will contain most of the

injected NH3, the concentration may exceed acceptance

limits for ash salability. Additional treatment of fly ash

holding ponds and basins may be required if large

amounts (30 ppmv) of SO3 are being removed. The use

of NH3 for SO3 mitigation is practical on units equipped

with SCRs, because NH3 is used as the reagent and is

readily available on site.

Fuel Switching and Blending

The firing of sub-bituminous coals typically results in low

SO3 formation and emission rates. Consequently, switching from firing bituminous to sub-bituminous coals can

be a mitigation option. However, equipment and fuel cost

factors often make such a change impossible. Many boiler

systems do not have the capacity and the equipment to

accommodate the firing of sub-bituminous coal without

major modifications that make fuel switch economically

unacceptable. Coal availability and costs also can constrain fuel switching. One possible solution is the blending of bituminous and sub-bituminous coals to create a

blend that has fuel and ash characteristics favorable for

SO3 emissions. This strategy is currently used for controlling SO2 emissions. When sub-bituminous coals are

blended with bituminous coals, the overall S content of

the fuel is reduced, resulting in a reduction of the SO2

concentration in the flue gas as well as the conversion of

SO2 to SO3 in the furnace and SCR and, thereby, resulting

in an overall reduction in SO3. Also, the sub-bituminous

coal ash contains large percentages of alkaline materials

that further assist in the capture of SO3 in the APH and

ESP.

Wet ESPs

Similar to dry ESPs, wet ESPs (WESPs) operate in a threestep process involving (1) charging of the entering particles; (2) collection of the particles on the surface of an

oppositely charged surface; and (3) cleaning the collection surface. Both technologies employ separate charging

and collection systems. However, the collecting surface in

WESPs is cleaned with a liquid, in contrast to mechanical

cleaning in dry ESPs. Consequently, the two technologies

differ in the nature of particles that can be removed,

the overall efficiency of removal, and the design and

maintenance parameters.47 While dry ESPs are typically

limited to power levels of 100–500 W per 1000 cfm of flue

gas, WESPs can operate with power levels as high as 2000

W per 1000 cfm. Because of wet cleansing of the collection system, PM does not accumulate in the collection

electrodes; this mitigates particle re-entrainment. Based

on these factors, WESPs can collect sub-micron particles

and acid mist very efficiently. WESPs can be configured

for vertical or horizontal gas flows in tubular or plate

designs. Tubular designs offer smaller footprints and, in

general, are more efficient than the plate type.

WESPs can be integrated easily with a wet scrubber.

In fact, integration of the WESP within the wet scrubber is

a design option with many attractive features including a

compact footprint; the ability to integrate the handling of

the wash water and solids from the WESP with scrubber

slurry, thereby avoiding the need for separate tank and

blowdown system; and the ability to collect the fine

H2SO4 mist, which typically escapes the scrubber because

of its very small droplet size.47

In 1986, the first commercial WESP application on a

U.S. power plant took place when AES Deepwater, a

155-MW cogeneration plant firing petroleum coke as the

primary fuel, was equipped with a WESP. With the WESP

in operation, the plume opacity at the plant is generally

10% or less.48

Recently, an upflow tubular design WESP has been

retrofit at Northern States Power Company’s Sherco Station in a wet scrubber/WESP configuration. Two more

power plant applications are underway presently: (1) a

5000-cfm slipstream at Bruce Mansfield Station; and (2) a

plate-type WESP for integration with Powerspan’s ECO

technology to be demonstrated at First Energy’s R.E.

Burger plant.

Tests at the Sherco Station (WESP retrofit to the outlet

section of the wet scrubber) allowed the scrubber to maintain a 70% SO2 reduction while keeping particulate emissions at 0.01 lb/106 Btu and opacity under 10%. Full

conversion of all scrubber modules at the plant with

WESPs is now underway.49 The WESP at the Mansfield

Station is achieving greater than 95% removal of SO3 and

PM2.5 and stack flow with near 0 opacity.50

Changing the Operation of APH

Increasing the heat transfer (or cooling of the flue gas) in

the APH would appear to be a potentially viable strategy

for removing some of the SO3/H2SO4 in the APH. This, in

turn, would lead to increased condensation of H2SO4 in

the APH and also to improved plant efficiency. However,

the dew point at the APH outlet is 230 °F (110 °C),

thereby limiting the SO3 removal to 90%.30 On the

other hand, the potential for corrosion in the APH and

downstream duct would increase with increased H2SO4

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condensation. Consequently, more frequent soot blowing

may be required to control this corrosion. These factors

would need to be considered when deciding to change the

APH operation to mitigate SO3 emissions. Data reflect that

25% increase in H2SO4 condensation may be possible by

lowering the flue gas temperature at the APH exit by 40

°F (22 °C).

SUMMARY

Formation and emissions of SO3/H2SO4 potentially can

lead to plant operation-related problems and environmental concerns. The formation of SO3 is complex, depending upon fuel, operating parameters, and plant configuration, and understanding the parameters leading to

excessive generation of SO3 and subsequent formation of

H2SO4 can assist in selections of practical control approaches. Elevated SO3 concentrations can lead to corrosion, the formation of sulfite scale, and fouling and plugging of low-temperature plant components, and can add

to the particle loading to control equipment. In some

cases, elevated SO3 concentrations can lead to the formation of visible plumes at the stack exit or shortly downstream of the stack, resulting in noncompliance with local

regulations.

SO3 can be formed in the boiler during the combustion of S-bearing fuels or downstream, particularly in SCR

reactors. SO3/H2SO4 can condense on low-temperature

components and gets adsorbed by fly ash. Adsorption is

much greater in the high alkaline ashes resulting from

sub-bituminous coal firing. SO3/H2SO4 traditionally has

been measured using extractive controlled condensation

methods.

If needed, SO3/H2SO4 emission can be mitigated using a variety of methods. Injection of alkali materials into

the furnace, either with the fuel or in slurry form, has

resulted in reductions of up to 80%. Post-furnace injection of alkali materials can achieve up to 90% reductions

but can increase particle loadings and ash resistivity characteristics. NH3 injection can also reduce SO3/H2SO4 by

90% and may result in increased particle loading to the

downstream collection systems. In plants with adequate

operational and equipment flexibility, fuel switching and

blending can be used to reduce formation and emissions

of SO3. WESPs are also an option for control of SO3/

H2SO4, and a variety of designs have been successfully

demonstrated for collection of acid mists and opacity

control.

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About the Authors

R.K. Srivastava and C.A. Miller are at the Office of Research

and Development, National Risk Management Research

Laboratory, Research Triangle Park, NC. C. Erickson and R.

Jambhekar are at Babcock Borsig Power, Inc., Worcester,

MA. Address correspondence to: R.K. Srivastava, Office of

Research and Development, National Risk Management

Research Laboratory, Air Pollution Prevention and Control

Division, U.S. Environmental Protection Agency, Research

Triangle Park, NC, 27711; fax: (919) 541-0554; e-mail:

srivastava.ravi@epa.gov.

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